Fluxes of nitrogen oxides above a subtropical forest canopy in China

Science of the Total Environment 715 (2020) 136993

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Fluxes of nitrogen oxides above a subtropical forest canopy in China Piaopiao Ke a, Qian Yu a, Yao Luo a, Ronghua Kang b,⁎, Lei Duan a,⁎ a b

State Key Laboratory of Environmental Simulation and Pollution Control, School of Environment, Tsinghua University, Beijing 100084, China CAS Key Laboratory of Forest Ecology and Management, Institute of Applied Ecology, Chinese Academy of Sciences, Shenyang 110016, China

H I G H L I G H T S

G R A P H I C A L

A B S T R A C T

• NOx flux above a subtropical forest was observed by aerodynamic gradient method. • A net dry deposition of 6.27 kg N ha−1 a−1 occurred as the net NOx exchange rate. • The atmospheric NOx concentration was the most important controlling factor. • More than 40% of deposited NOx might be uptake by the canopy.

a r t i c l e

i n f o

Article history: Received 13 October 2019 Received in revised form 12 January 2020 Accepted 27 January 2020 Available online 28 January 2020 Editor: Jianmin Chen Keywords: Aerodynamic gradient method NOx flux Dry deposition Subtropical forest Canopy uptake

a b s t r a c t Dry deposition of Nitrogen (N) in forests is commonly estimated from inferential method and/or throughfall measurements, with inevitable uncertainty. In this study, we applied an aerodynamic gradient method to directly measure the nitrogen oxides (NOx) flux above the canopy of a subtropical forest in southeastern China for two consecutive years. The flux and transfer velocity generally reached the maximum absolute values in the midday, with the largest diurnal maximum of absolute flux values observed in the winter of 2015 and that of transfer velocity in the autumn of 2015. The annual average transfer velocity was −0.79 and −0.38 cm s−1 in 2015 and 2016, respectively. Although the net downward NOx fluxes predominated for both years, upward flux (net emission) of NOx was observed during spring months, which reflected the possible bi-directional exchange balanced by soil-atmosphere and foliage-atmosphere exchanges. The NOx concentration seemed to be the most important factor controlling the NOx exchange above canopy, and could mainly explain the seasonal variation of N deposition. The linear regression between the NOx flux and concertation was explored, and it was observed that the deposition of NOx was offset by possible underlayer emission of NOx when the ambient NOx concentration below1.7 ppbv and 1.9 ppbv at night and in the day, respectively. The average dry deposition of NOx for the two years was 6.28 ± 0.06 kg N ha−1 a−1, N40% of which might be uptake by the canopy, estimated by comparing the wet/ throughfall deposition measurement of nitrate with the observation of NOx flux. This indicated the importance of stomatal uptake of NOx in nitrogen budget in subtropical forests. © 2020 Published by Elsevier B.V.

1. Introduction Nitrogen oxides (NOx, mainly as NO and NO2) are important air pollutants as being precursors in the formation of fine particles (Seinfeld and Pandis, 2012), photochemical smog (Altshuller and Bufalini, ⁎ Corresponding authors. E-mail addresses: [email protected] (R. Kang), [email protected] (L. Duan).

https://doi.org/10.1016/j.scitotenv.2020.136993 0048-9697/© 2020 Published by Elsevier B.V.

1971), and acid deposition (Duan et al., 2016). The NOx produced from anthropogenic or natural sources (Galloway et al., 2004) are subsequently oxidized to HNO3 or nitrate (NO− 3 ) in the atmosphere, with the ultimate fate of being scavenged down in the aqueous form as wet deposition or being directly transported to the terrestrial surface in the gaseous or particulate form as dry deposition (Wolff et al., 2010). The enhanced nitrogen (N) deposition (both wet deposition and dry deposition) to the terrestrial ecosystems, if not exceeding their tolerant

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capacity, could improve their productivity, and increase carbon uptake and sequestration (Liu and Greaver, 2010; Hietz et al., 2011; Min et al., 2014). However, the excess N input could lead to N saturation, which is characterized by significant NO− 3 leaching and possibly results in severe acidification of soil, eutrophication of water, and reduction in plant growth and ecosystem biodiversity (Dentener et al., 2006; Xie et al., 2018). It is crucial to understand N deposition in the forest comprehensively, considering the great importance of forest in the biogeochemical cycles of N. Wet deposition of N has been widely monitored in the world (Vet et al., 2014; Duan et al., 2016), whereas there is a large scarcity of direct measurement of dry deposition, primarily due to the sophisticated instrumentation and methods of the observation (Vet et al., 2014). Dry deposition of N consists of oxidized N (NOy) and reduced N (NHx), where the former mainly involves NOx, HNO3, HONO, aerosol nitrate, and organic particle molecules (Geddes and Martin, 2017). Generally, NOx deposition composes 12–50% of the total dry deposition of NO y (Zhang et al., 2005; Zhang et al., 2009; Xu et al., 2015; Jia et al., 2016; Zhao et al., 2017; Yu et al., 2019), with further dominance at highly NOx polluted sites (Zhang et al., 2005). As an easier substitution of direct observation of dry deposition to forests, throughfall measurement is widely used, and the difference between throughfall and wet deposition is acknowledged as an approximation of dry deposition to the forest ecosystems (Duan et al., 2016). However, in the case of NO x, the dry deposition process is much more complicated, as leaf-level uptake of NOx has been validated in the lab experiment and field investigation (Chaparro-Suarez et al., 2011; Breuninger et al., 2013; Delaria et al., 2018), and this part of dry deposition could be neglected by throughfall method (Chen and Mulder, 2007; Wright et al., 2016; Thimonier et al., 2019). Furthermore, NO could be released from soil on the forest floor, which is subsequently transformed to NO 2 in the air and could account for the upward flux of NOx above forests (Neirynck et al., 2007; Seok et al., 2013; Min et al., 2014). Due to very limited field measurements, the dry deposition of NOx to the forests is generally estimated through the inferential technique in most studies (Flechard et al., 2011; Kharol et al., 2018). An inferential model determines the flux of NOx by multiplying the ambient NOx concentration with modeling-deduced deposition velocity (Shen et al., 2009), and may thus have large uncertainties (Wesely and Hicks, 2000; Flechard et al., 2011). In recent years, the applications of critical load in policydecisions (Duan et al., 2015; Zhao et al., 2017) require accurate estimations of total N deposition to calculate critical load exceedance (the N deposition in excess of the ecosystem critical load). Long-term, direct micrometeorological measurement of reactive nitrogen flux is thus strongly recommended for validation of inferential models (Flechard et al., 2011). As an important part of NOy dry deposition, NOx flux above the forest canopy have been seldom measured in temperate forests in Europe (Foken et al., 2012) and North America (Horii et al., 2004; Seok et al., 2013; Geddes and Murphy, 2014; Min et al., 2014), and in Amazon rainforest (Rummel et al., 2002; Andreae et al., 2015). In contrast, there are quite few direct observations of NO x exchange above subtropical forests, which have been increasingly identified as hotspot of nitrogen deposition (Geddes and Martin, 2017; Schwede et al., 2018), particularly in southern China (with high N deposition averaged to 23 kg N ha −1 yr−1 ) (Liu et al., 2013). In this study, the flux of NO x above a subtropical forest canopy was continuously monitored for two consecutive years in southeastern China, and the diurnal, monthly and seasonal variation of flux were also presented. In addition, the flux determined by the direct observation was compared with the measurement of throughfall and wet deposition of nitrate, and the stomatal uptake of NOx was then estimated.

2. Materials and methods 2.1. Site description The flux and concentration of NOx were measured on a flux tower from March 2015 to February 2017, at Qianyanzhou (QYZ) station (115°4′E, 26°45′N), which is managed by the Chinese Academy of Sciences (CAS). The site, surrounded by Pinus massoniana forest (N600 m to the forest edge, around 17 ha in area, with average canopy height of 16 m; (Yu et al., 2018b)), and farmland outside the forest, is about 42 and 15 km from the nearest city (Ji'an) and county (Taihe), respectively (Fig. 1). Both the city and county, where many anthropogenic NOx emission sources such as coal-fired boilers and automobiles exist, are located to the northwest of the site, while a large mountain area without obvious anthropogenic sources is to the southeast. The wind that blows from the northwest direction brings significantly higher concentration of NOx than from the southeastern direction (Fig. 2). The atmospheric N deposition at this site was reported to be 44.8 kg N ha−1 a−1 via throughfall measurement (Cheng et al., 2017). QYZ is under typical subtropical monsoon climate, with an annual average temperature of 17.9 °C and annual average precipitation of 1494 mm during 1985–2005 (Xu et al., 2011). 2.2. Monitoring method and flux calculation Generally, there are three approaches to monitor the surface exchange flux of NOx: chamber measurement, aerodynamic gradient method, and eddy covariance method. The chamber measurement is often applied for soil emission (Butterbach-Bahl et al., 2004; Kang et al., 2017) and plant-atmosphere exchange (Chaparro-Suarez et al., 2011; Breuninger et al., 2012). Suitable for measurement of gas exchange between the atmosphere and the whole ecosystem, the eddy covariance method has been applied to measure NO and/or NO2 flux in several field investigations recently (Min et al., 2014; Lee et al., 2015; Vaughan et al., 2016). However, the eddy covariance measurement requires fast-response analysers detecting over 5 Hz, which could limit the use of this method. Moreover, most of NOx flux observations above canopy using this method only lasted several weeks or a growing season (Rummel et al., 2002; Horii et al., 2004; Stella et al., 2013; Geddes and Murphy, 2014; Min et al., 2014), which was not long enough to complete the budget of dry deposition of NOx and possibly due to the challenges in maintenance (Wu et al., 2015). Since the dry deposition of NOx occurs with temporal variation (Horii et al., 2006; Zhang et al., 2013), a long-term evaluation, covering the four seasons in a year is necessary. Compared with the eddy covariance method, the aerodynamic gradient method may be more suitable for long-term observation of NOx flux above forest canopies (Stella et al., 2012), and was thus applied in this study. The aerodynamic gradient method used to calculate the flux of a trace gas is based on the theory that the flux is related with the turbulent condition and vertical concentration gradients, being expressed as follows: FC ¼ KC KC ¼

∂c ∂z

u κz Φc ðζ Þ

∂c C −C h2 ðh1 Nh2 Þ ≈ − h1 h1 −h2 ∂z Vd ¼

FC C h1

ð1Þ ð2Þ

ð3Þ

ð4Þ

where FC is the vertical trace gas flux, KC is the turbulent exchange

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Fig. 1. The location of QYZ sites, the nearby county (Taihe County) and city (Ji'an City) (Google Earth). Vegetation map of China as background (CAS, 2007).

coefficient (or K value), u∗ is the friction velocity, κ is the Karman constant (0.4), z is the measurement height, Φc(ζ) is the integrated universal stability function of concentration when the atmospheric stability is ζ, ∂c/∂z is the vertical NOx concentration gradient, Ch1 and Ch2 are NOx concentrations at height of h1 and h2 respectively, and vd is the transfer velocity. In this study, both K value and concentration gradient between 25 m and 35 m were calculated every 5 min and further averaged to 1 h. All the metrological sensors (Table S1, apart from the sensor in the soil, including total solar radiation, air temperature, humidity, wind direction, and wind speed), the sample inlets, and pipelines were set at 25 m and 35 m of the flux tower, separately, with the NOx analyser in a cabin near the tower (Fig. S1). The data of the meteorological sensors were stored as five minutes' mean on the data logger (CR 1000, Campbell Scientific Inc., USA). The air intakes and sensors in this study were settled outside the tower body N1 m away to avoid the disturbance from the tower. The K value was calculated based on the atmospheric stability correction (Businger et al., 1971; Yu and Sun, 2006), following certain steps written in MATLAB (2017b, Mathworks). Detailed description of the calculation was shown in Section S1 of SI. The atmospheric concentration of NO and NO 2 was simultaneously measured with a chemiluminescence analyser (Model 42i Trace Level, Thermo Environmental Instruments, USA). The solenoid switched every 5 min, controlled by a time relay, allowing

gas flowing alternatively from two heights into the analyser. The pipeline, air intakes, and valves were all made of Teflon to avoid the adsorption of NO x. The particle filter in the front of the pipe was replaced every week, and the time relay was adjusted twice a week to match the time of NOx analyser. Considering only 30% of NO concentration gradient was higher than the detection limit and the inter-cycling reaction between NO-O3-NO2 does not influence the total amount of NOx (sum of NO and NO2), only the fluxes of NOx were discussed in this article. 2.3. Data quality control The NOx analyser was calibrated by standard NO-contained gas every month. The concentration data due to the malfunction of the NOx analyser, such as low convert temperature inside the analyser, were rejected. The flux data were also screened for outliers, with values outside the monthly mean ± 3 standard deviations being rejected. In this study, the data with insignificant concentration gradient were retained since the flux was often very small under this condition and the removal of these data would possibly lead to an overestimation of the average flux (Converse et al., 2010). For NOx, complex chemical reactions with O3, hydroxyl radical (OH), and organic nitrates, has been reported within the forest canopy (Min et al., 2014). Thus, the chemical reaction time scales (τchem) and

Fig. 2. Seasonal wind frequency in 2015 (a) and 2016 (b). The length of the wedge is the frequency of wind in the season and the color indicates the concentration of NOx.

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turbulent time scales (τturb) should be compared before the application of gradient method in calculating the turbulent transport (Neirynck et al., 2007; Mayer et al., 2011). The analysis of τchem and τturb revealed that the chemical removal time of NOx at the observation height was generally much larger than the turbulent transport time (Section S2 and Fig. S2 of SI). Thus, the gradient method was applicable for monitoring the NOx flux at this site. The overall uncertainty of the flux calculation was about 4.5% (detailed description in Section S3 of SI) and the valid data coverage was 88% and 80% in 2015 and 2016, respectively (Table S2).

Vd (cm/s) Concentration(ppbv) Gradient(ppbv/m)K value(m2/s)

(a)

10

Spring Summer Autumn Winter

3. Results and discussions 3.1. Diurnal, monthly and seasonal variation of NOx fluxes The annual flux of NOx was −29.9 ng N m−2 s−1 (−32.1 to −27.6 ng N m−2 s−1, 95% confidence interval) and −9.77 ng N m−2 s−1 (−11.0 to −8.59 ng N m−2 s−1, 95% confidence interval) in 2015 and 2016, respectively. The annual mean transfer velocity was −0.79 cm s−1 (−0.84 to −0.73 cm s−1, 95% confidence interval) and −0.38 cm s−1 (−0.47 to −0.29 cm s−1, 95% confidence interval). This

(b) Spring

Summer Autumn Winter

5

0

0.10 0.05

Flux (ngN/m2/s)

0.00

10

5

0

1 0 -1 -2 -3 -4

0.2 0.0 -0.2

0

8

16

0

8 16 Autumn

0

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0 -50

3 2 1 0 -1 -2 -3 0

-100

8 16 Winter

-150 0 8 16 0 8 16 0 8 16 0 8 16 Hour of day (Local)

0 8 16 0 8 16 0 8 16 0 8 16

Hour of day (Local)

Fig. 3. Diurnal variations of K value, NOx concentration gradient, mean NOx concentration for two heights, the transfer velocity and NOx flux in 2015 (a) and 2016 (b). The data was averaged from the hourly data in each season. The shadow area indicates the standard error. The negative values of flux mean deposition to the forest while positive values mean the emission from the forest.

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in 2015 and the spring in 2016. The absolute value of transfer velocity of NOx above the canopy calculated in this study was larger than previously summarized 0.2 cm s−1 from the flux monitoring over a temperate deciduous forest (Horii et al., 2004). One thing to note is that the transfer velocity here was calculated from the net flux and the concentration, which is not the same as the traditional theoretical deposition velocity. Theoretically, the flux of NOx above the forest canopy is determined by the relative importance of soil-atmosphere exchange and foliageatmosphere exchange, as well as chemical reaction between NOx and free radical groups (Wesely and Hicks, 2000; Ganzeveld et al., 2002; Hari et al., 2003; Foken et al., 2012). Here, NOx (sum of NO and NO2) as a whole was considered a quasi-stable gas, because the specific time for vertical mixing has been much shorter than chemical reaction (Stella et al., 2013) (Fig. S2). Also, the photolysis of nitrate on the leaf surface has been suggested as sources of NOx emission under strong UV-light (N10 W m−2, Hari et al., 2003). The foliage-atmosphere NOx exchange was recently reported to result in N deposition even when the ambient concentration of NOx (9 m above the canopy) under 1 ppbv (Chaparro-Suarez et al., 2011; Breuninger et al., 2013; Delaria et al., 2018). Thus, the NOx could be assumed to be uptake by the foliage at this site as the ambient NOx concentration was much larger than 1 ppbv almost all the time. It is generally acknowledged that leaves could absorb NO2 through stomata and eventually assimilate it into amino pools (Chaparro-Suarez et al., 2011; Delaria et al., 2018). This procedure is mainly positively controlled by the stomatal aperture and ambient concentration (Chaparro-Suarez et al., 2011). The stomatal uptake of NO2 in the daytime might enhance the transfer velocity as double of the night (Horii et al., 2004). The transfer velocity has also been proved to be proportional to the stomatal conductance (Horii et al., 2004; Breuninger et al., 2013). The vertical mixing was also faster during the day, which could be seen by larger K value (Fig. 3), and also

denoted the net NOx deposition in both years. Diurnal variation of NOx fluxes was clearly shown in Fig. 3. In 2015, similar to observations of NOx fluxes above temperate forests (Horii et al., 2004; Geddes and Murphy, 2014; Min et al., 2014), the largest absolute values of both flux and transfer velocity during one day were usually observed at midday. At night time, the transfer velocities and fluxes were close to zero in the spring (from March to May) and summer (from June to August), and remained negative in the autumn (from September to November) and winter (from December to February). The maximum midday absolute values of flux in spring, summer, autumn, and winter of 2015 were 8.36 ng N m−2 s−1, 10.9 ng N m−2 s−1, 66.2 ng N m−2 s−1, and 139 ng N m−2 s−1, respectively, with positive value observed only in the spring. This indicated occasional NOx emission from the forest in the spring in the midday and net NOx deposition in the other three seasons. In 2016, the daily variation of NOx flux in the spring and summer was similar to that in the winter of 2015, with a midday NOx deposition up to 82.8 ng N m−2 s−1 and 8.55 ng N m−2 s−1, respectively. In the autumn and winter of 2016, the maximum absolute value was close to 1.85 ng N m−2 s−1 and only fluctuated in a small range (Table S3). The nighttime (19:00–07:00) deposition was generally smaller than that at the daytime (07:00–19:00) (Fig. 3). The monthly NOx flux in 2015 showed a small net emission from March to May (with the maximum emission in April, 13.7 ng N m−2 s−1), and net deposition starting from June and continuing increasing to the maximum of 145 ng N m−2 s−1 in December (Fig. 4). In 2016, the monthly NOx deposition declined from March to June, and thereafter remained smaller. Among these seasons, NOx deposition occurred in the order of: winter N autumn N summer N spring (emission in the spring in 2015), and spring N summer N winter ≈ autumn in 2016. Similar to variation of diurnal pattern of fluxes in different seasons, the transfer velocity varied in larger scales and displayed a relatively large absolute value (around 1.00 cm s−1) in the autumn and winter

(a)

500 2015 rainfall 2016 rainfall

30 20

300

10 200 0 100

-10

0

-20

(b) Monthly average flux of NOx (ng N/m2/s)

400

40 0 32 -100

24 flux, 2015 night flux, 2015 day con, 2015 night con, 2015 day

-200

flux, 2016 night flux, 2016 day

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con,2016 night con, 2016 day

8

Concentartion (ppbv)

°

2015 temperature 2016 temperature

Rainfall (mm)

Temperature ( C)

40

0

-300 Mar

Apr

May

Jun

Jul

Aug

Sep

Oct

Nov

Dec

Jan

Feb

Month Fig. 4. The monthly averaged temperature and rainfall (a); NOx flux and NOx concentration (abbreviated as con in the legend) (b).

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enhanced the deposition of NOx. The variation of concentration in different seasons might induce the dramatically decrease of negative fluxes in the autumn and winter in 2016 (see more detailed description in Section 3.2). In the daytime of the spring in 2015, the upward flux of NOx from the forest canopy (3.0–8.0 ng m−2 s−1, 95% confidence interval) was observed, while strong deposition was observed in the spring of 2016. Upward flux of NOx has also been observed in temperate forests, and was interpreted as a result of the forest soil emission of NO (Neirynck et al., 2007; Flechard et al., 2011). The NOx generated from soil could either be uptake by plants (Min et al., 2014), or induce an upward flux above the forest canopy as long as being large enough (Flechard et al., 2011). The observed NOx concentration on the forest floor of this site in 2015 and 2018 (ranging 3.18–5.12 ppbv in several days of the September in 2015, and 7.96–8.39 ppbv in the summer and autumn of 2018, 95% confidence interval, unpublished data) was far smaller than derived soil-atmosphere compensation point (a concentration below which soil would act as a source instead of a sink of NOx) of forest soil in other studies, ranging from 9 ppbv to 500 ppbv (Godde and Conrad, 2000; Bargsten et al., 2010; Behrendt et al., 2014). In addition, the subtropical forest soil, under chronically high N deposition, was suggested as a pronounced NO emission source (13.4–24.5 ng N m−2 s−1) in the summer of southwest China (Kang et al., 2017). Meanwhile, the higher concentration of NO at the lower height in the midday of the spring in 2015 also indicated strong upward emission of NO from the forest (Fig. S4). Thus the NO emission from the forest floor might have the potential to invert the canopy NOx flux from deposition to emission, particularly in the season when the NOx concentration was relatively low. The field investigation of soil NO emission has also found that higher temperature could induce stronger soil emission, while increased rainfall (accompanying with lager air humidity and soil water content) is likely to dampen the soil NOx emission (Butterbach-Bahl et al., 2004; Kang et al., 2017). Comparing the results in the spring of these two years, a larger deposition was monitored in 2016 (Fig. 4). In 2016, the average temperature, 20.5 °C, was higher than that in 2015, 18.6 °C, and the precipitation events occurred more in the spring and summer of 2016, whereas the precipitation was evenly distributed in the four seasons of 2015 (Fig. S3). The rainfall was significantly larger in the spring of 2016, accompanying with higher soil water content (Fig. S4) and this could inhibit the soil emission of NO (Butterbach-Bahl et al., 2004; Kang et al., 2017). Moreover, the more rainfall and higher temperature in the spring of 2016 might enhance the growth of needles, which could lead to larger LAI and thus increase the foliage uptake (Flechard et al., 2011). Besides, similar to the flux, the seasonal transfer velocity in the two years displayed a dramatic variation. As described above, the NOx flux observed above the forest canopy was the net flux with downward leaf deposition of NO2 abated by the upward soil NO emission. When the ambient NOx concentration was high, the soil NO emission might be inhibited (Bargsten et al., 2010), while the leaf NO2 deposition was enhanced (Chaparro-Suarez et al., 2011). Under this context, the transfer velocity might reach large values as observed in the autumn and winter of 2015. Correspondingly, the soil NO emission in the spring of 2016 was estimated to be smaller than that in 2015, with leaf NO2 deposition varying on the contrary. This might induce largely increased negative transfer velocity in the spring of 2015.

coincided with the declining trend of NOx emission in this province (0.54, 0.49, 0.41 and 0.36 million tons in 2014, 2015, 2016 and 2017, respectively; National Bureau of Statistics of China, 2019). Since 2013, China has implemented the Clean Air Action plan to reduce national emissions of major air pollutants including NOx (Zheng et al., 2018). Since the emissions from coal-fired power plants, with much higher chimney and stronger regional affects than other sources such as vehicles (Lu et al., 2016), were preferentially controlled, the NOx emission from nearby power plants was possibly greatly reduced, and eventually led to the distinct decrease of NOx concentration in remote areas such as at this site. The monthly NOx concentrations in 2015 showed a similar trend to the flux, decreasing from March to May, remaining at around 2ppbv in September, and then increasing to 9 ppbv (the maximum in the year) in December (Fig. 4). In 2016, the NOx concentration decreased from March to June and remained between 2 and 4 ppbv after September, much smaller than the corresponding concentration in 2015, accompanying with a significantly smaller deposition. This further emphasized the important role of NOx concentration in deciding the flux at this site. The NOx flux was most strongly correlated with the NOx concentration, other than the meteorological parameters, e.g. the temperature and relative humidity in both years, particularly in the seasons with relatively high NOx concentration (Table S4). When the ambient NOx concentration was below 1.9 ppbv and 1.7 ppbv in the daytime and nighttime respectively (Fig. 5), the net flux was positive, and the NOx deposition was likely to be offset by the underlayer emission processes, mostly the soil NO emission and possible surface chemical reactions. Furthermore, when extrapolating the ambient NOx concentration to zero, positive NOx fluxes were expected and estimated to be around 21.1 ng N m−2 s−1 and 28.8 ng N m−2 s−1 in the daytime and nighttime, respectively. This might indicate the magnitude of the NOx emission from the subtropical forest at this site. In addition, the R-square in the winter and autumn was much larger than observed in the spring and summer, mainly driven by the large deposition at high NOx concentration (N5 ppbv). As reported above the temperate Harvard forest (Horii et al., 2004), the correlation coefficient of multi-regression between the flux and concentration became less statistically significant and the R-square decreased under low NOx concentration compared with high NOx concentration (N2 ppbv). Different from the small soil emission of NO in Horii et al. (2004), the soil NO emission had the potential to invert the direction of NOx flux above forest canopy in this study as described above. Under low NOx concentration, the soil emission of NO possibly becomes comparable with the deposition above the forest canopy, as the latter is positively related with the NOx concentration in theory (Wesely and Hicks, 2000) and validated in this study and other field investigation (Horii et al., 2004; Geddes and Murphy, 2014). Precisely, the canopy surface flux of NOx was considered as the net exchange rate of soil-atmosphere and foliage-atmosphere (Wesely and Hicks, 2000; Ganzeveld et al., 2002). Although the relation between NOx concentration and the corresponding flux may thus not be fully explained by the linear regression, that the NOx flux scaled with the NOx concentration at the study site clearly underlined the vital importance of NOx concertation to the flux. 3.3. Estimation of canopy uptake by comparing the direct monitoring and throughfall measurement

3.2. Dominant role of NOx concentration in determining the exchange flux The NOx deposition above forest canopy was significantly, positively correlated with the atmospheric NOx concentrations (Fig. 5 and Table S4). The annual atmospheric concentrations of NOx above the forest canopy were 4.53 ppbv (4.44–4.62 ppbv, 95% confidence interval) and 2.32 ppbv (2.28–2.36 ppbv, 95% confidence interval) in 2015 and 2016, respectively. A significant decrease of NOx concentration in the autumn and winter of year 2016 has been monitored (Fig. 4). This

In this study, the annual flux of NOx above the canopy was monitored to be downward, which yielded a net dry deposition of NOx 6.27 ± 0.06 kg N ha−1 a−1 averaged for the years of 2015 and 2016. The dry deposition of NOx at QYZ site was much higher than that directly observed above a managed ponderosa pine plantation in the east America (below 2.5 kg N ha−1 a−1; Horii et al., 2004; Min et al., 2014) and a coniferous forest in the eastern Canada (below 1 kg N ha−1 a−1; Geddes and Murphy, 2014), ranging also much higher than

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Fig. 5. The linear regression between the flux and concentration of NOx during the night (a) and the day (b). The flux and concentration data were hourly 20-days averaged.

estimated annual NO2 deposition in southeastern China (below 5 kg N ha−1 a−1; Xu et al., 2015; Zhang et al., 2017) from inferential method and chemical transport models. Still, it was in the upper range of NO2 deposition to the forest ecosystems modelled across Europe (up to 9.4 kg N ha−1 a−1), when the ambient NO concentration was high (7–8 ppbv) (Flechard et al., 2011) and the site was exposed to anthropogenic pollution (Neirynck et al., 2007). At the same site, the total deposition of NO− 3 estimated from throughfall measurement (TF) in 2014 (Cheng et al., 2017) and 2017 with the same method (unpublished data) were 16.3 ± 0.9 and 8.1 ± 0.4 kg N ha−1 (Fig. 6), respectively. Simultaneously, the nitrate wet deposition (WD) were 14.0 ± 1.0 and 3.2 ± 0.2 kg N ha−1, respectively. The estimated dry deposition of oxidized nitrogen (TF-WD) was thus much lower than the dry deposition of NOx (DDNOx) (Fig. 6), which implied the canopy uptake of NOx possibly larger than 2 kg N ha−1 a−1. In other word, the canopy uptake of NOx could contribute as much as 40% of the total dry deposition of NOx. Although there were inevitable uncertainties while comparing the estimated TF-WD with DDNOx in different years, it was still reasonable to assume the average throughfall and wet deposition in 2015 and 2016 as the mean value of those in year 2014 and 2017 respectively, when considering the evidently declining trend of both the NOx emission in this province and NO− 3 -N deposition at this site. Similarly, the differences found between total NOy dry and fog deposition estimates and NO− 3 net throughfall fluxes suggested that

approximately 50% (6.4 kg N ha−1 a−1) of the total NOy deposition was irreversibly retained within the canopy of European forest (Erisman and de Vries, 2000). The foliar uptake of NOx has also been estimated through stomatal conductance evaluation and ambient concentration measurement, ranging 1.2–3.1 kg N ha−1 a−1 in subtropical forests (Hu et al., 2016) and temperate forest (Garcia-Gomez et al., 2018). The relatively larger NOx concentration (Garcia-Gomez et al., 2018) and unsaturated forests of N (Skeffington and Hill, 2012; Yu et al., 2018a) might lead to the relatively larger stomatal uptake of NOx at this site. Considering the possible soil emission of NO, the retention of nitrogen oxides in the canopy might be even higher. Meanwhile, in the winter and autumn, the dry deposition of NOx far overweighed the dry deposition of other oxidized nitrogen species (indicated by TF-WD ≪ DDNOx). In Masson pine dominated subtropical forest, there was no significant difference of LAI in four seasons (Zeng et al., 2008). As described above, the stomatal uptake of NOx is mainly impacted by ambient NOx concentration and stomatal aperture (Chaparro-Suarez et al., 2011; Breuninger et al., 2012). Even though the stomatal resistance in the growing season are estimated to be the minimum (Zhang et al., 2017), the NOx concentration in the autumn and winter was much higher than other seasons, and possibly stimulated the stomatal uptake of NOx. It implied that the stomatal uptake of NOx in the subtropical forest should take an important role in the total nitrogen budget in the forest. The absorption of NOx by canopy was also widely observed in the

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20 WD of 2014 WD of 2017

18

DDNOx TF-WD (2 year average) WD (2 year average)

16 Deposition (kg N ha-1a-1)

concentration was much lower in 2016 than in 2015, which mainly led to significantly decrease in NOx dry deposition. In recent years, the anthropogenic emission of nitrogen oxides is gradually reduced in China (Yu et al., 2019), which results in a decreased ambient air NOx concentration. The oxidized N deposition, particularly dry deposition, is foreseen to decline (Yu et al., 2019), accompanying with decreasing NOx concentration, as observed in this study. Detailed mechanism controlling the dry deposition process and future changes of the dry deposition require further investigation.

TF of 2014 TF of 2017

14 12 10 8 6

Declaration of competing interest

4 2 0 Spring

Summer

Autumn

Winter

Annual

Fig. 6. Comparison between nitrate‑nitrogen throughfall deposition (TF), wet deposition (WD), and monitored net deposition of NOx (DDNOx). Mathematically, when the value, TF-WD-DDNOx, became negative, the canopy uptake was larger than the absolute value of TF–WD-DDNOx, and the lower limit of canopy uptake of NOx was illustrated as – (TFWD-DDNOx).

throughfall measurement in temperate forests (Erisman and de Vries, 2000; Neirynck et al., 2007). For subtropical forest ecosystems, previous studies using the throughfall measurement could possibly underestimate the oxidized nitrogen deposition to forests (Fang et al., 2011; Du, 2018). 4. Conclusion The aerodynamic gradient method has been applied in two consecutive years to investigate the NOx fluxes above a subtropical forest in southeastern China. Due to the stable and practicable characteristics of the gradient-flux observation system, a high data coverage (N80%) was attained, which enabled calculation of total deposition/emission of NOx to/from the forest ecosystem. The diurnal variation of fluxes in four seasons showed that the downward fluxes dominated, and stomatal uptake of NOx was likely to provoke stronger deposition in the day. Similar to the diurnal variation of fluxes, the NOx transfer velocity displayed the maximum values in the day apart from the autumn and winter in 2016, when the fluxes became quite small. Furthermore, the fluxes and transfer velocity also varied seasonally, with larger absolute values in the autumn and winter of 2015 and the spring of 2016. The transfer velocity of NOx in this study was likely to be related with the NOx concentration. The annual mean transfer velocity was −0.79 ± 0.03 cm s−1 and −0.38 ± 0.05 cm s−1 in 2015 and 2016, respectively. In the spring of 2015, an upward flux of NOx was observed, while the maximum absolute values of fluxes were often observed in the day and being negative in other seasons. Compared between the springs in two years, the drier soil in 2015 might lead to stronger soil NO emission and eventually induce upward fluxes of NOx above the forest canopy. Furthermore, the NOx flux at this site was predominantly controlled by the NOx concentration. The linear regression between the flux and concentration indicated that the leaf uptake of NO2 was likely to be offset by the soil NO emission when the ambient NOx concentration was below 1.7 ppbv and 1.9 ppbv during the night and day, respectively. Meanwhile, the correlation between NOx concentration and the flux was decreasing with the declined NOx concentration. The relatively more important role of soil NO emission under low ambient NOx concertation could be responsible for this phenomenon. Additionally, the overall dry deposition of NOx (6.27 ± 0.06 kg N ha−1 −1 a ) was compared with the throughfall measurement of nitrate in a previous study at this site. The results suggested that N2 kg N ha−1 a−1, or as much as 40% of the NOx was possibly be uptake and permanently retained in the canopy. It should also be noted that the NOx

The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. Acknowledgement The authors are grateful for the financial support of the special fund of State Key Joint Laboratory of Environmental Simulation and Pollution Control (grant number 19L02ESPC) and the National Natural Science Foundation of China (grant number 41877329 and 21607019). The authors also greatly acknowledge the support from Qianyanzhou Forest Experimental Station and the help received in system maintenance from Yuanfen Huang. Appendix A. Supplementary data Supplementary information. Sections S1–S3: calculation steps of turbulent transfer coefficient, the characteristic time of chemical reaction and turbulent transport. Tables S1–S3: the seasonal data coverage, seasonal atmospheric concentration, flux and meteorological parameters, and correlation coefficient between flux and other parameters. Figs. S1–S4: a sketch of the monitoring system set-up, the characteristic time of chemical reaction and turbulent transport, hourly variation of meteorological parameters and flux measurement, and diel variations of NO and NO2 concentration. Supplementary data to this article can be found online at https://doi.org/10.1016/j.scitotenv.2020.136993.

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